A Review of Flood-Related Storage and Remobilization of Heavy Metal Pollutants in River Systems
Water Air Soil Pollut
A Review of Flood-Related Storage and Remobilization of Heavy Metal Pollutants in River Systems
Dariusz Ciszewski 0 1
Tomáš Matys Grygar 0 1
0 T. M. Grygar Faculty of Science, J.E. Purkyně University in Ústí nad Labem , Ústí nad Labem , Czech Republic
1 T. M. Grygar Institute of Inorganic Chemistry , AS CR, v.v.i., Řež , Czech Republic
2 ) AGH University of Sciences and Technology , Krakow , Poland
Recently observed rapid climate changes have focused the attention of researchers and river managers on the possible effects of increased flooding frequency on the mobilization and redistribution of historical pollutants within some river systems. This text summarizes regularities in the flood-related transport, channel-to-floodplain transfer, and storage and remobilization of heavy metals, which are the most persistent environmental pollutants in river systems. Metaldispersal processes are essentially much more variable in alluvia than in soils of non-inundated areas due to the effects of flood-sediment sorting and the mixing of pollutants with grains of different origins in a catchment, resulting in changes of one to two orders of magnitude in metal content over distances of centimetres. Furthermore, metal remobilization can be more intensive in alluvia than in soils as a result of bank erosion, prolonged floodplain inundation associated with reducing conditions alternating with oxygen-driven processes of dry periods and frequent water-table fluctuations, which affect the distribution of metals at low-lying strata. Moreover, metal storage and remobilization are controlled by river channelization, but their influence depends on the period and extent of the engineering works. Generally, artificial structures such as groynes, dams or cut-off channels performed before pollution periods favour the entrapment of polluted sediments, whereas the floodplains of lined river channels that adjust to new, post-channelization hydraulic conditions become a permanent sink for fine polluted sediments, which accumulate solely during overbank flows. Metal mobilization in such floodplains takes place only by slow leaching, and their sediments, which accrete at a moderate rate, are the best archives of the catchment pollution with heavy metals.
River; Sediment; Heavy metals; Mobilization; Pollution; Flood
Economic development, which has rapidly grown since
the Industrial Revolution, has been accompanied by an
increasing demand for heavy metals and substances
containing metal compounds. Heavy metals escape
during ore extraction and processing and are also
widespread in industrial and municipal sewages; their
sources have been extensively reviewed in earlier works
(e.g. Förstner and Wittmann 1981; Salomons and
Förstner 1984). For almost 200 years of extensive metal
utilization, the possible toxic effects of the intake of
heavy metals were not recognized, and industrialization
has resulted in their uncontrolled dispersal in hundreds
of kilometres of river systems.
Heavy metals are discharged in both dissolved and
solid phases in proportions that vary greatly depending
on the element properties, pollution sources and
chemistry of receiving river waters. In river water, metals tend
to precipitate rapidly or to be adsorbed onto solid
particles. These processes may be reversed with changes in
the Eh and pH, which are master variables that control
the partitioning of metals between sediments and the
water column (Salomons and Förstner 1984). In river
systems, concentrations of the same metal in
finegrained sediment can be one to several orders of
magnitude higher than in the dissolved phase (Martin and
Meybeck 1979; Horowitz 1991). Freshly deposited
sediments may easily liberate metals during resuspension
with an increase of flow velocity (Salomons et al. 1987).
A fine fraction of sediment, comprising silt and clay, has
been recognized as the crucial component of the
pollution load stored within many river systems (Owens et al.
2005), whereas heavy metals associated with
coarsegrained sediments constitute an important part of the
pollution load over short reaches of some mine-affected
rivers (Marron 1989; Ciszewski 1998). Conventionally,
the part of metals, which passes through 0.45-μm filter,
is named Bdissolved^, although it does not represent
truly dissolved metal ions. It is composed of free metals,
complexions or metals bounds to ligands, which may
aggregate into organic or inorganic entities of colloidal
size (1.2 μm–1 kDa). Metals associated with colloids
cannot settle themselves until aggregated into larger
particles and can be conveyed with dissolved phase over
long distances (Gueguen and Dominik 2003).
Sediment-associated heavy metals can be stored in
fluvial systems for periods from days to millennia,
depending on the river-flow dynamics. It has been
estimated that river channels of mid-sized catchments, of
the order of thousands square kilometres, are
responsible for the storage of only a small percentage of the
annual sediment-associated pollutant flux (Walling et al.
2003; Villarroel et al. 2006). The storage of fine-grained
sediment in channelized urban rivers is of the order of
days to months (Taylor and Owens 2009). In channels of
less modified rivers, sediments of fine-grained sand and
silt fractions can be stored for several to tens of years in
shelters, at confluences or in wide channel sections as
sediment Bplugs^ and sequences many decimetres thick
(Skalak and Pizzuto 2010; Faměra et al. 2013). For most
perennial river systems, overbank sedimentation is
considered to represent longer-term storage for fine
sediments with a much larger residence time of the order
of 102–103 years (Matys Grygar et al. 2016a),
representing a net loss to actual downstream sediment
conveyance (Walling et al. 1999). The floodplain
pollutant flux depends on sediment deposition, which is
typically a few tens of percentage points of annual
sediment load delivered to the main channel system. It
depends on the actual sediment budget; e.g., in the
midsized catchment of the River Aire, approximately one
third of historical anthropogenic Pb has been stored in
the floodplain (Walling et al. 2003).
Floods are known both for their devastating potential
of human infrastructure and for maintaining valuable
riparian communities. They also play a crucial role in
creating and reshaping the dispersal pollutant patterns in
a river system (Matys Grygar et al. 2014). During
floods, the pollutants formerly temporarily stored in
the channel are quickly entrained and transferred to the
floodplain (Matys Grygar et al. 2014). In turn, the
erosion and leaching of the polluted floodplain
sediments was recognized as important factor influencing
secondary river pollution proportional to the degree of
historical pollution and dilution by the enhanced input
of particulates from the watershed and bank erosion
(Navrátil et al. 2008; Chen et al. 2014). Floods may also
trigger primary pollution when precipitation extremes
cause failures of settling ponds or washes from
stockpiles. A special case of an artificially enhanced
discharge-driven pattern is a Bpollution pulse^—a
sediment wave or sediment slug—the introduction of extra
material, e.g. due to mining operations (Miller et al.
1998), which is then only slowly further transported
downstream by the trunk river.
Given that previous reviews have considered impact
of metal mining on the aquatic environment (Byrne et al.
2012; Wolkersdorfer 2004; Miller 1997) or essentially
have summarized the effects of mining on the rate of
natural fluvial processes and contamination along
mineaffected rivers (Macklin 1996), this review focus on
artificially modified rivers, which have been usually
contaminated from numerous point sources. Such rivers
are widespread in densely populated and industrialized
areas of the world and characterized by fluvial processes
altered due to engineering structures. Particular attention
is given to floods and we stress that floods and
highwater stages are the reasons, which speed up metal
circulation in river valleys by orders of magnitude if
compared to the non-inundated soils. Unlike previous
reviews (Du Laing et al. 2009; Schultz-Zunkel and
Krueger 2009), we do not deal with estuaries with
salinity effects on metal mobility and give more
attention to fluvial processes as they have principal effects on
metal dispersal. We review the effects of floods and
high-water stages on the dispersal of heavy metals in
channels and floodplains in seven main sections. We
start with the description of the hydraulic control of
heavy-metal storage in a channel (Section 2), modes of
channel-to-floodplain heavy-metal transfer (Section 3),
the longitudinal and spatial patterns of floodplain
storage (Section 4), the influence of channel engineering on
heavy-metal storage (Section 5), the role of
impoundments in the storage of heavy metals (Section 6), the
influence of floods on metal remobilization (Section 7)
and methods of pollution mapping (Section 8).
2 Flood Control on Heavy Metals’ Storage
Floods play the most important role in the transport of
heavy-metal pollutants associated with particulate
matter, particularly in severely polluted catchments. In such
river systems, both concentrations of suspended
particulate matter and pollutant contents increase with the
growing discharge, particularly in the early stage of
floods. Their values sometimes remain high even during
the flood attenuation (Müller and Wessels 1995;
Baborowski et al. 2004; Coynel et al. 2007a, b;
Resongles et al. 2015), producing a counter-clockwise
hysteresis loop of total pollutant loads versus discharge
(Zonta et al. 2005; Coynel et al. 2007a, b). The loop
documents the activation of pollutants and sediments as
enhanced runoff and river discharge pass through their
temporary sink. Higher discharges may also increase
dissolved pollutant concentrations, e.g., in the case of
As; by contrast, the dissolved concentrations of most
other pollutants, i.e., those primarily transported in
particulate forms, are decreased due to dilution by excess
water (Baborowski et al. 2004; Resongles et al. 2015).
It was found that periods of low flows in dry seasons
lead to higher concentrations of heavy metals in
channel-bed sediments, whereas wet seasons are
characterized by a lower metal content in the bed and a
higher metal content in suspended sediments (Gaiero
et al. 1997; He et al. 1997). Physical seasonal changes
may be exacerbated by changing redox conditions at the
river bottom, associated with varying organic matter and
Fe- and Mn-oxide contents in sediments (Gaiero et al.
1997; He et al. 1997). Floods may also change the
riverwater chemistry. In acidic waters, such as those
containing acid-mine drainage, the enhanced input of rainwater
may dilute excess acidity and promote the hydrolysis of
Fe3+ ions, followed by the precipitation of Fe oxides and
sweeping a part of dissolved heavy metals to solid
particles (Cánovas et al. 2012).
Floods shift the boundary between bedload and
suspended load by re-suspending the fine sediment
fractions from their temporary sinks in the channel.
The storage of sediment-associated heavy metals in
a channel is a complex interplay of erosion,
accumulation and sediment reworking. In a given cross
section, the intensity of these processes depends on
the turbulence and velocity of current flow, which
is a function of the channel shape and the bed
morphology producing remarkable spatial
heterogeneity. In a sand-bed alluvial, perennial river, the
current flow intensity is reflected in metal content
in a fraction <1 mm (Ciszewski 1998). Metal
content in fine silt and clay, <0.063-mm fraction,
reflects the metal concentration observed in a
suspended load. The fine-grained fraction
accumulates primarily in the shelters near channel banks,
where the flow velocity enables the settling of
small and low-density particles (Rhoads and Cahill
1999). A part of fine-grained sediments is also
trapped from suspension by plants and plant roots
producing homogeneous fine-grained sediments,
best suitable for river monitoring (Ciszewski 1998;
Nováková et al. 2013). The lowest heavy-metal
concentrations usually occur in channel bars;
however, on a gravel bed river, polluted fines can
infiltrate in greater amounts into gravel substrate
during low-water stages (Ladd et al. 1998).
Furthermore, being the deepest places in the channel,
pools can be a place to store fine sediments during
low-water stages and for the precipitation of
manganese and iron hydroxides, with some metals, on
gravel surfaces (Evans and Davies 1994). During
floods, sediment-associated heavy metals are
scoured from pools (Ciszewski 1997). On the
contrary, in channels of episodic rivers, the maximal
concentrations occur in the zone of the most
frequent flow, and much smaller concentrations are in
the near-bank deposits (Graf et al. 1991). The
complexity of the channel dynamics and the transient
nature of channel sediments make their stratigraphy
too complicated to reconstruct the pollution history
(Nováková et al. 2013; Faměra et al. 2013).
In river reaches with permanently active pollution
sources, the effect of flushing of polluted, fine-grained
sediments by floods is limited, and metal concentrations
can quickly return to the previous state. On the Biała
Przemsza River, concentrations of zinc, lead and
cadmium dropped approximately 3- to 4-fold following a
100-year flood over a 40-km reach downstream the
discharge point of effluents from the lead–zinc ore mine
in southern Poland (Ciszewski 2001). The drop was
accompanied by the coarsening of near-bank channel
deposits. In the reach adjacent to the pollution source,
over the year following the flood, concentrations of
metals increased, whereas downstream reaches
exhibited further decreases in concentrations. On reaches at
former mine sites, overbank sediments contribute to
the pollution of the channel, delaying pollution decrease
by natural dilution with extra clean material. In some
cases, it is estimated that natural decay of metals within
a river channel may last several hundred years (Moore
and Langner 2012). On the Carson River in Nevada,
pollution of the channel sediment was not markedly
altered even by 100-year flood due to the mobilization
of polluted overbank sediments (Miller et al. 1999). The
long-term stability of channel-sediment pollution is also
observed on small perennial streams despite numerous
floods passing through a river system. Investigations of
the Matylda stream sediments pollution by heavy metals
indicated only a small decrease over 40 years after
mining cessation (Ciszewski et al. 2012).
Nevertheless, the heavy-metal dispersal pattern in
channel sediments can change in particular
crosssections after flood flows, and concentrations can
drop, go up and stay the same in particular channel
locations (Protasowicki et al. 1999; Moody et al.
2000). On the Mała Panew River, incised in sandy
alluvia, pollutant (Ba and Zn) concentrations were
observed for 2 years in the same channel locations
(Ciszewski 2004). In these places, the migration of
sandbars several metres long and several tens of
decimetres high resulted from the passing of two
floods of moderate magnitudes. Usually, a dead
zone, characterized by flow velocities below
<0.1 m/s, appeared in front of the bar at low
discharges. It occurred for a few months until it was
eroded or filled with the sand of the prograding
dune. As a result, changes in the flow velocity were
observed in the same places of the channel. These
changes were well correlated with changes in the
organic matter content and in the heavy-metal
concentrations (Fig. 1).
3 Modes of River-to-Floodplain Heavy-Metal
Heavy metals are transferred from a channel to the
floodplain surface only when water table exceeds the
bankfull stage. Because approximately 90 % or more of
metal load can be associated with sediment particles, the
pathways of metals to the floodplain are essentially the
same as those of suspended sediments (Wyżga and
Ciszewski 2010). Dissolved heavy metals are
considered to play a minor role in the transfer of metals to the
floodplain (Hostache et al. 2014). However, according
to some estimates, even if the volume of water infiltrated
into the floodplain were a few percentage points, it
would correspond to approximately 10 % of the mass
of pollutants deposited on the floodplain per a single
flood event (Stewart et al. 1998). Although the
prediction of infiltrated contaminants has been neglected in
most chemical mass-balance studies, conditions that
favour adsorption onto sediments can affect their
entrapment (Gonzalez-Sanchis et al. 2015).
When flood water overtops the river bank, the
contrast between the flow velocity in the channel and that in
the floodplain produces eddies, which transfer solid
particles and momentum from the deeper and faster flow
in the channel to the shallower and slower flow over the
floodplain (Knight and Shiono 1996). This mechanism,
described by the diffusion mixing model, results in levee
deposits, thickness and grain size of which decrease as
the water-flow velocity slows down with an increasing
distance from the channel (Pizzuto 1987). The other
mechanism of sediment transfer by convection may
occur where there is a component of flow perpendicular
to the channel. Convection may result in strong
differences in the amount of particles accumulated across the
floodplain because the flow of water can transfer coarse
particles on the floodplain by tractive movement
(Marriott 1992). Sediment diffusion in lowland alluvial
rivers results in the highest accumulation rate of metal
load immediately at the river bank, whereas their
maximum concentrations occur in thinner strata outside the
levee zone if metal pollutants are primarily associated
with finer size fractions. The sediment transfer by
convection occurs in crevasse splays associated with a less
Fig. 1 Changes of heavy metals concentrations, losses on ignition, content of fine fraction and water depth in the same place of the channel
are related to channel bar migration during the 2 years period (from Ciszewski 2004, modified)
uniform distribution of metal load across the floodplain
and the highest metal concentrations behind the zone of
coarse-grained deposits (Wyżga and Ciszewski 2010).
Convective sediment transport across the floodplain
is particularly effective on sinuous rivers. River
sinuosity induces helicoidal currents, which are responsible for
the flood deposition of sandy fractions on the convex
banks. For this reason, the thickness of sandy deposits
can be higher on sinuous sections than on straight
reaches modified by channel-training works (Ten
Brinke et al. 1998). The strength of the helicoidal
currents increases as discharge rises, and large floods
produce lateral accretion deposits during point-bar
formation in the near bank zone by both diffusion and
convection processes (Hooke and Le 1975). These
processes are active on natural rivers because stream-bank
reinforcements reduce channel shifting on channelized
4 Floodplain Storage of Heavy-Metal Pollutants
The large variability of heavy-metal distribution in
floodplains outlines the heterogeneous nature of
sediment deposition. The maximal pollution may be
concentrated in nearly isolated hotspots, discontinuous
zones or strata, which decline markedly over distances
of metres (Heaven et al. 2000; Ciszewski et al. 2008;
Matys Grygar et al. 2014, 2016a). The most polluted
sediments can be found in point-bar deposits of shifting
channels and in natural levees when heavy metals are
associated primarily with ore grains of high density
(Marron 1989) or when they contain finer sediment
laminae (Matys Grygar et al. 2016a). On rivers, where
heavy-metal pollution is primarily associated with the
finest size fractions, floods may create a thin but
spatially extensive blanket of polluted overbank fines.
Considerable pollution may also be stored in abandoned
meanders (oxbow lakes) formed by cut-offs just before
or during periods of pollution climax (Matys Grygar et
Generally, the role of floods in the redistribution of
polluted sediments within a river system is relative to its
magnitude, and smaller river systems preserve the
primary effects of large historical pollution for longer time.
A closer look at the general trend of the fast longitudinal
decrease in pollution usually shows a more complicated
picture (Bird et al. 2008) and cannot be interpreted
without a detailed and geomorphic description of
individual rivers. The maximal heavy-metal concentrations
may be lower in a floodplain than in a channel, but their
decrease with the growing distance from the pollution
source is usually less steep (Heaven et al. 2000; Macklin
et al. 2003).
In formerly mined regions, a considerable portion of
river systems was modified by extra sediment load
inserted in the climax of mining. The associated
heavy-metal concentrations could now be flattened and
smeared by flood-induced reworking or preserved in
terraces. Floodplain aggradation, usually enhanced
during the mining period, was followed by erosion after
mining cessation. In such cases, maximal pollution
could be found in sediment bodies elevated above the
current active floodplain (Brewer and Taylor 1997;
Macklin et al. 1994; Miller et al. 1998). With time,
persistent sediment reworking by floods decreases the
extent of pollution with a growing distance from the
former mines, and sometimes the current pollution
maxima are further downstream from the original
sources (Miller et al. 1998; Dennis et al. 2009; Foulds
et al. 2014).
Particularly serious and spatially extensive floodplain
pollution is caused by tailing dam failures or slurry
remobilization, usually triggered by precipitation
extremes (Hudson-Edwards 2003; Hudson-Edwards et
al. 2001; Bird et al. 2008; Žák et al. 2009; Matys Grygar
et al. 2014). Because such pollution events are
associated with floods to which huge volumes of severely
polluted slurry are inserted, these Bpollution pulses^
may produce widespread pollution peaks in sediment
records (Resongles et al. 2014; Matys Grygar et al.
2014, 2016a), unless the river system has been severely
polluted before the event (Hudson-Edwards 2003;
Hudson-Edwards et al. 2001; Bird et al. 2008). The
event layer can also be identified by different pollutant
ratios (Resongles et al. 2014) or a specific isotope
signature (Matys Grygar et al. 2014, 2016a).
If the floodplain part outside of a levee zone is
regularly inundated by overbank flows, and if the river
transports a sufficient amount of clay and silt fractions to
produce yearly increments strata of at least several
millimetres, sediment profiles with a stratigraphic order
can be sampled there and then used as pollution archives
(Grosbois et al. 2012; Nováková et al. 2013; Van Metre
and Horowitz 2013; Dhivert et al. 2015b). The effect of
sediment sorting on element concentrations must be taken
into account to distinguish temporal and grain-size
controls (Dung et al. 2013; Chen et al. 2014; Bábek et al.
2015). Geochemical normalization is more efficient for
this purpose than conventionally used sieving (Kersten
and Smedes 2002). Normalization considerably improves
extraction of historical pollution signal from lithological
variability inherent to fluvial systems and impacting also
overbank fines. Sediment dating can be performed by
means of the gamma spectrometry of fallout
radionuclides or by a correlation of the rapid growth of metal
concentrations or metal peaks in a series of several
profiles with production characteristics known from the
industrial history of a drainage basin (Ciszewski and Malik
2004; Lokas et al. 2010). Coarser (sandy) intercalations
of extreme flood layers may also be used for their dating
(Dhivert et al. 2015b; Zhang et al. 2015). In floodplains, it
is necessary to distinguish overbank sequences from
lateral channel deposits, the latter being less suitable for
dating because of the contrasting rate and style of their
deposition (Matys Grygar et al. 2013). The advantages of
the sequences of overbank fines from the distal part of the
floodplain as sedimentary archives are the relatively
uniform stratigraphy, a lithology that prevents the vertical
migration of pollutants in top strata, and a lower
probability of erosion gaps (Lewin and Macklin 2003). The
disadvantages of overbank sediment archives are the
dilution of pollutants at overbank discharges (Nováková
et al. 2013; Matys Grygar et al. 2013) and the risk of
postdepositional processes due to reductimorphic processes
or the physical translocation at depths closer to the
groundwater table (Ciszewski et al. 2008; Du Laing et
al. 2009). Furthermore, oxbow lake sediments may be
valuable sedimentary archives. In them, the time of
meander cut-off, either artificial (Van Metre and Horowitz
2013; Sedláček et al. 2016) or natural (Matys Grygar et
al. 2016a), can be recorded as a lithological change; it
provides a valuable time constraint for the younger
sediments. Their deposition rate can be of the order of cm/y
(Sedláček et al. 2016).
5 Effect of Channel Engineering on Heavy-Metal
Channel straightening by artificially cutting off river
meanders is the most widespread modification of the
river channel used to improve navigation and to obtain
land for agriculture. Channel cut-offs are artificial
features of the floodplain landscape resulting from
intentional channel straightening. However, channel
straightening may also occur naturally and produce oxbow
lakes, particularly in river valleys with a low gradient.
Artificial paleomeanders usually occur close to the
channelized river and act during floods as efficient traps
for sediment-associated heavy metals and other
contaminants. The proximity of the river channel favours the
rapid flood-related filling of cut-off channel segments,
e.g. with 2-m deposits within less than 150 years,
providing detailed records of the river pollution (Gocht et
al. 2001). The filling processes are rapid, accurately
reflecting changes in the river-sediment chemistry, as
long as the channel cut-off is connected with the river
(Bábek et al. 2008). When sediment is trapped only
during flood episodes, the pollution record and metal
concentrations are affected by the magnitude and
frequency of overbank flows and the flow regulation and
bed degradation following channelization (Dhivert et al.
2015a). Additionally, ponds may sometimes form on
embanked river floodplains located in sites of historic
dike breaches. These scars on the lower Rhine
floodplain trapped as much as 6 m of fine,
metalpolluted sediments over recent two centuries
Channel training works associated with bank lining
or groynes induces lateral channel stabilization, and
channel narrowing, even by several times, expands the
floodplain (Ciszewski and Czajka 2015). These
processes usually disturb the natural regime of sediment
transport because channel straightening causes an increase in
the channel gradient and the rapid erosion of the
riverbed and banks in the period immediately following the
channelization works (Łajczak 2003). In channels cut in
fine-grained alluvia, the rapid erosion of river banks can
even led to increases in the channel width and the
formation of a braided channel pattern (Kiss and Sipos
2007). A high rate of sediment accumulation on such an
artificially shrunk floodplain is usually associated with
channel degradation during channelization and can be
even several times higher than before the
channeltraining period (Kiss et al. 2011). After the period of
channel adjustment to new hydraulic conditions, the rate
of overbank sand transfer onto the floodplain decreased,
whereas finer grained sandy deposits with abundant clay
layers became typical floodplain-accretion deposits, as
observed on the Rhine River (Hesselink et al. 2003).
Lateral channel fixation by bank reinforcements, which
usually follow river straightening, confines the zone of
lateral channel sediment accretion to the very narrow
strip of land along channel banks, leading to the
progressive channel narrowing and deepening. In lined and
laterally stabilized channels, overbank deposition
becomes the dominant process in the development of the
floodplain (Ciszewski and Czajka 2015). Moreover, as
the main floodplain-forming process, overbank
deposition is also constricted by flood dykes to the narrow strip
of the original floodplain, leading to a continuous
increase in the floodplain elevation and the inherent, on
many floodplains, decrease in the flooding frequency
and sediment-accumulation rate (Hobo et al. 2014).
Groynes have been routinely constructed at the
riverbanks of channelized European lowland alluvial rivers
since the nineteenth century to direct the flow current to
the channel centre and thus to enhance bed scour and to
improve navigation conditions. Usually, groynes are
directed at small angles against the flow direction to
enable sediment entrapment (Sukhodolov et al. 2002).
Groyne basins are known as sinks for fine, polluted
sediments during low and average water stages, whereas
during floods, they are sources of pollution (Baborowski
et al. 2012), which is favoured by low compaction and
large sediment thickness (Schwartz 2006). However, for
example, in the historically contaminated Odra River,
the groyne basins became long-term sinks for
contaminated sediments because they operated as particularly
effective sediment traps. With time, as the surfaces of
the infill grew during high discharges, they were
progressively keyed into the floodplain, resulting in a
twoto threefold reduction in the channel width (Ciszewski
and Turner 2009). Currently, groyne deposits in the
incised, upper river reaches form approximately
4-mthick sequences of fine sediments contaminated with
heavy metals. The sediments contain black layers of
coal particles intercalated with clean sands. The zone
of these laminated deposits, which accumulate with a
high average rate of approximately 5 cm/year, is
confined to the width of the pre-regulation channel and does
not exceed approximately 30 m on average (Ciszewski
and Czajka 2015). In the middle river, reach stabilized
by regulation structures, sediments contaminated with
heavy metals occur in three zones of variable widths
(Fig. 2). The zone of the nineteenth century groyne
basins filled with nineteenth and twentieth century
fine-grained sediments ranges in width from 10 to
approximately 100 m, and the thickness of these deposits
reaches 3 m. The most intensive polluted sediment
accretion, with a rate of 5 cm/year in a 40-year period,
has taken place in a river reach heavily polluted by
mining and urban effluents (Ciszewski and Czajka
2015). This rate is among the maximum values observed
for human-modified rivers (Provansal et al. 2010).
Generally, river channelization with groyne
construction changes natural sediment cycling in a river valley,
and in a polluted river, it produces a heavy-metal
dispersal pattern different from that in non-engineered
channels. This difference is reach specific and depends
on the width of the pre-regulation channel, the length of
groynes, the dynamics of the channelized river, the
suspended load transported, the degree of river pollution
and the length of the heavy pollution period. Heavy
metal-contaminated deposits of the peak of the
industrial era can be stratified. They contain a high
organicmatter content, and commonly, parts of refuse material,
e.g., bricks, plastic and ashes, can be called industrial
alluvium (Ciszewski and Czajka 2015; Lewin 2013) or
agro-industrial alluvium (Macklin et al. 2014). These
deposits can also be characterized by a lack of
bioturbation and upward fining or coarsening, and they
usually accrete with the highest rate in groyne basins or in
Fig. 2 Distribution of metal-contaminated sediments in zones along the channelized reach of the Odra River is related to the width of the
pre-regulations channel and repeated channel training works in 19th and 20th century (from Ciszewski and Turner 2009)
near-bank shelters other than those on the associated
floodplain (Swennen and Van der Sluys 2002). They
represent a novel, artificial sedimentary facies.
River embankments are perhaps the most widespread
modifications in the river valleys of densely populated
areas. Embankments confine flood-inundation zones to
the relatively narrow strip of the original floodplain and
reduce the area of overbank sediment accretion. These
modifications are the most influential in river reaches
conveying large amounts of sediment. In the upper
Vistula River reach with the high sediment-accretion
rate, the inter-embankment floodplain level is as much
as 2 m higher above the surface outside embankments
(Łajczak 1995). Rapid sediment deposition on the upper
Vistula in the twentieth century was coeval with the
peak of pollution, leading to the accumulation of thick
sequences of sediments strongly polluted with heavy
metals. Pollution with metals is two orders of magnitude
lower in sediments of pre-industrial era behind flood
dykes where zinc, cadmium and lead levels are close to
the local geochemical background (Macklin and Klimek
1992). In most floodplains of lowland rivers draining
industrialized areas, the average rate of sediment
accretion decreases from approximately 1 cm in the
proximity of the river channel to a few millimetres per year at
flood dykes (Asselman and Middelkoop 1995; Kiss et
al. 2011). For this reason, the thickness of the polluted
sediments of the industrial era in the inter-embankment
zone usually does not exceed several decimetres
(Middelkoop 2002; Overesch et al. 2007; Ciszewski
2003). Generally, dykes seem to diminish the retention
of pollutants by the shortening inundation time rather
than to favour their conveyance losses because most
dykes cut off depressions of the backswamp zone,
where prolonged water stagnation with fine sediment
6 Storage of Heavy Metals in Dam Reservoirs
Dam reservoirs are traps for solid particulates (Van
Metre and Horowitz 2013) and associated heavy metals
(Palanques et al. 2014) transported through the river
system. Globally, as much as 20–30 % of sediment
transported by rivers is trapped in reservoirs, but this
amount varies locally, primarily in relation to the
reservoir depth, catchment topography and land use of the
catchment (Vörösmarty et al. 2003; Syvitski et al. 2005).
Reservoirs trap extremely variable parts of the annual
metals load, which can reach 90 % for Pb, Cd and Cu
(Schintu et al. 1991). Most deposits are accumulated in
the shallow backwater zone, but the most polluted
deposits usually occur in the deepest sections of reservoirs,
where the undisturbed settling of the fine-grained clay
and organic sediments takes place (Zhao et al. 2013).
The deep parts of reservoirs where thermal stratification
develops are characterized by oxygen depletion
eventually leading to anoxic conditions, which triggers the
reduction of nitrate, Fe and Mn hydroxides and sulphate
(Friedl and Wüest 2002). Anoxic conditions in bottom
sediments also cause the reduction and methylation of
Hg, which enhances its export both as dissolved MeHg
and bound in small organisms to remarkable distances
downstream (Schetagne et al. 2009; Carrasco et al.
2011). Enhanced MeHg production immediately after
the impoundment concerns all reservoirs, natural and
artificial; however, the problem is particularly appealing
when the sediments have already been polluted by Hg
(Carrasco et al. 2011). Audry et al. (2010) described the
conversion of labile Zn phases, silicates and oxides to
sulphides, sulphates and Fe-oxide-associated species in
the reservoir bottom. The reservoir sediment sinks are,
however, not permanent: Changes in redox conditions
(oxidation), acidification or physical disturbances,
including flood inflows, can liberate the pollutants back
to the water column (Coynel et al. 2007a, b; Audry et al.
2010; Yang et al. 2014; Hamzeh et al. 2014). At a large
discharge or water release from the reservoirs, the
unconsolidated sediments are easily re-suspended and
turned back to the river system (Bi et al. 2014).
Floods and high river discharges may result in
seasonal changes in heavy-metal concentrations in
reservoir-bottom sediments, which document the
transient nature of reservoir storage. Large floods, which
flush sediment-associated heavy metals from polluted
catchments, may increase the content of some metals in
waters and sediments in the summer or during
springmelting periods (Kwapuliński et al. 1991;
SzarekGwiazda et al. 2011). In the following part of a year,
metals may be remobilized from the bottom and exit the
dam lake (Kocharyan et al. 2003). In some years, fluxes
of metals, mobilized from reservoir sediments, can be
higher than metal loads entering that reservoir (Rzętała
2008). The enhanced heavy-metal pollution of bottom
sediments is also observed in shallow reservoirs of
heavily industrialized regions during dust emission in
the cold season of the year (Reczyńska-Dutka 1985).
Dams tend to decrease the variation of fluvial
discharges and to supress flooding and overbank
deposition. Damming and discharge regulations prevent
deposition in elevated surfaces in floodplains between
extreme floods (Dhivert et al. 2015b) and makes
deposition in low-lying surfaces more regular. Due to the
decreased fluvial transport of solids, the pollution
downstream from the dams may be less diluted by upstream,
usually cleaner sediments (Van Metre and Horowitz
Dams change a spatial pattern of fluvial sediments
exposed to the variation of redox conditions. The
change in the ratios of risk elements, i.e., Pb and Zn
on one hand and Fe on the other, in littoral areas
subjected to regular (seasonal) water-level changes in a dam
have recently been described (Liu et al. 2014). It is
probable that prolonged waterlogging near the reservoir
shore with a variable level will promote redox-driven
processes and enhance mobility in a periodically or
permanently inundated littoral zone.
The depth profiles in dam-reservoir sediments may
be particularly valuable pollution archives (Audry et al.
2004; Sedláček et al. 2013). The sedimentary
environment is much less variable in reservoirs than in
floodplains and has much less reworking in deep, quiet
locations. Their permanent existence under the water
column, which decreases post-depositional migration,
may provide more detailed reconstruction of historical
pollution (Matys Grygar et al. 2012). Seismic profiles
(Palanques et al. 2014) and other geophysical
techniques (Bábek et al. 2008) are efficient in distinguishing
reservoir sediments from pre-dam deposits. The onset of
the reservoir deposition is a robust date point; otherwise,
dam sediments can conveniently be dated by the gamma
spectrometry of fallout radionuclides 210Pb and 137Cs
(Sedláček et al. 2013, 2016). Other date points can be
obtained by the assignment of coarser (sandier)
intercalated sediments to extreme floods (Bábek et al. 2011;
Dhivert et al. 2015b).
7 Flood-Related Remobilization of Heavy Metals
Historically contaminated floodplains constitute a
secondary source of river pollution. Its significance
depends on the extent of physical and chemical processes
operating on variable spatial and temporal scales
(Macklin 1996). Proportions of the actual mechanisms
at play are usually not distinguishable. The existence of
many pollutant pathways varies depending on
individual pollutants, catchment characteristics and changes
with precipitation intensity and flood-wave discharge
(Zonta et al. 2005; Coynel et al. 2007a, b; Resongles
et al. 2015; Nováková et al. 2015).
Mobilization of sediment-associated heavy metals
from floodplains is controlled by erosion during flood
episodes, which on most of unregulated perennial rivers
occur every 2 years, on average (Petit and Pauquet 1997
and references therein). Riverbank erosion is most
intensive on natural, large meandering rivers. On
lowgradient alluvial plains, the lateral channel shifts reach
tens of metres per year, whereas on smaller rivers of
lower energy, they may be of the order of decimetres to a
few metres per year (Wang et al. 2014; Nicoll and
Hickin 2010; Black et al. 2010). The erosive supply of
the polluted sediment from riverbanks is reduced almost
to zero by channel lining or revetment and is
considerably suppressed by a variety of Bsofter^ engineering
measures, such as tree planting and channel dredging.
In these and other laterally stable river reaches,
floodplain erosion is limited almost solely to cultivated
surfaces and paleochannels or depressions where
highwater turbulence produces scars or chutes up to
hundreds of metres long (Navrátil et al. 2008).
In areas of former metal mining, large amounts of
pollutants may enter rivers through the flood-induced
undercutting of waste tips exposed directly on banks
(Foulds et al. 2014), the erosion of thick fine-grained
floodplain sediment sequences (Dennis et al. 2003; Žák
et al. 2009) or the erosion of paleochannels filled during
the mining era, produced by the cut-off and
abandonment of meander loops (Miller et al. 1998). In valleys
with sparse vegetation in dry climates, a single, large
flood may enlarge the river channel by two to three
times and result in channel shifting by over 100 m. In
such rivers, the erosion and redeposition of
sedimentassociated metal load is particularly effective along
reaches characterized by low gradients and wide valley
floors (Miller et al. 1999). Confined river reaches are net
erosional during floods due to bank erosion (Thompson
and Croke 2013) and act as transitional zones for
sediment–associated heavy metals and other contaminants
transported downstream (Macklin and Lewin 1989;
Graf 1990). Metals associated with fine-grained
sediments, stored in the alluvia of low-gradient reaches, are
preferentially washed away as suspended sediment,
even during small flood events, leaving behind coarser
sediment as bed material (Žák et al. 2009). In small,
severely polluted river reaches, metal loads in the
suspended matter transported during floods may be very
high, but their concentration always results from the
relative contribution of historically polluted sediments
and cleaner material eroded from cut banks, tributaries
and catchment surfaces. Typically, the highest floods
supply the largest loads of unpolluted sediments to the
channel, whereas metal concentrations in suspended
sediment tend to decrease with the increase in discharge
(Salomons and Eysink 1981; Hutchinson and Rothwell
2008; Žák et al. 2009; Schultz-Zunkel and Krueger
2009). The dilution by cleaner sediments may, however,
be insufficient to bring metal concentrations below
target limits downstream from severely polluted areas
(hotspots) where floods used to transport more polluted
sediments (Matys Grygar et al. 2014; Dhivert et al.
2015a). The actual concentrations of pollutants within
the flood wave are consequently a complex function of
discharge and sediment supply (Resongles et al. 2015;
Dhivert et al. 2015a). Due to different pollutant paths in
historically polluted fluvial systems, the actual ratios of
metal pollutants vary with the discharge (Resongles et
al. 2015; Nováková et al. 2015).
The chemical remobilization of metals during floods
is related to progressive oxygen depletion by microbial
and root respiration during floodplain inundation. With
the change from aerobic to reductive conditions, the
reductive dissolution of Fe and Mn hydroxides takes
place, and it is also controlled by pH, salinity, organic
matter content and temperature (Rinklebe and Du Laing
2011). In reducing conditions, metals such as Fe and Mn
and commonly associated pollutants such as As, Cd, Cr,
Mo, Ni and Zn can be released from the solid phase to
pore waters (Shaheen et al. 2014; Hindersmann and
Mansfeldt 2014). However, mobile fractions of metals
are not simply transported to water bodies, and metal
transfer is not currently quantifiable (Schultz-Zunkel
and Krueger 2009). Recently, diffusive gradient in thin
films (DGT) technique is used to determine pore water
profiles and remobilization of heavy metals at the
sediment/water interface (Wu et al. 2016). Flood
recession followed by the drying and aeration of floodplain
soils reverses the processes of metal dissolution. In an
oxic environment, Fe and Mn re-precipitate as oxides
and scavenge heavy metals back to the solid state (Du
Laing et al. 2009). This phenomenon is shown in depth
profiles of overbank sediments of the Ploucnice River in
the Czech Republic (Fig. 3). Iron and Pb are depleted in
reduced (grey coloured) strata and accumulated in
Feoxide rich concretions near boundary of reduced and
unaltered (brown coloured) overbank fines (Matys
Grygar et al. 2014, 2016a). Short-term flood episodes
do not affect pH and have lesser effect on metal
mobilization than redox changes. Moreover, it could be
expected that single and rapidly flowing flood waves have
a minor effect on the metal migration within floodplain
sediment profiles, but it was shown that in areas with a
longer flooding duration, the mobility of some metals,
expressed by their speciation, was higher than in areas
inundated less frequently (Shaheen and Rinklebe 2014).
Vertical metal displacement is favoured by the frequent
fluctuation of water-table levels and, in addition to the
visible accumulation of the secondary Fe and Mn
oxyhydroxides, may manifest in mineral breakdown
and pseudomorphing or high levels of exchangeable
and specifically adsorbed metals (Hudson-Edwards et
al. 1998). The consequences of past redox changes are
easily revealed as Berratic^ Fe and Mn depth profiles in
floodplain-sediment cores and systematic depletion at
depths below the common water table (Matys Grygar et
al. 2013). The translocation rate of metals can be high in
an acidic floodplain environment, and mobile metals
can retain in less acidic zones abundant in Fe oxides
(Kraus and Wiegand 2006), clay-rich zones (Cappuyns
Fig. 3 Depth profiles of Fe, U
and Pb in redox-altered and
polluted overbank sediments of
the Ploučnice River, the
Czech Republic; EFs (enrichment
factors) are concentrations
normalized to global (Fe) or local
(Pb and U) background values
(EF=1); Fe concentrations was
normalized by Al concentrations
and divided by mean upper
crustal global reference Fe/Al r
atio; U and Pb concentrations
were divided by local background
concentrations for the overbank
sediments of the studied river
(according to Matys Grygar et al.
2014; Matys Grygar et al. 2016a)
and Swennen 2004) or organic-rich zones, giving peaks
in vertical metal distribution (Ciszewski et al. 2008).
Metal redistribution with migrating groundwater has
been observed particularly in profiles of coarse-grained
deposits, which exhibit anomalous metal peaks at levels
related to the depth of the most frequent water-table
fluctuation (Taylor 1996; Ciszewski et al. 2008; Matys
Grygar et al. 2013). Although fluxes of heavy metals
with groundwater to a river can be extremely variable
locally and change over an order of magnitude at
distances of hundreds metres (Coynel et al. 2007a, b), their
role in natural systems is likely minor compared with
physical remobilization by fluvial erosion. Floodplains
as a diffusive source of heavy metals, are undoubtedly
most important during floods and high-water stages
(Aleksander-Kwaterczak and Ciszewski 2012;
Palumbo-Roe et al. 2012).
8 Mapping Floodplain Pollution by Heavy Metals
The spatial distribution of heavy-metal pollution in
floodplains has been documented by many authors
(Macklin et al. 1994; Miller et al. 1998; Heavens et al.
2000; Notebaert et al. 2011; Hobo et al. 2014; Foulds et
al. 2014). Investigations indicate that an extensive
sampling using depth profiles (sediment cores) is required to
faithfully describe it, and it is best if the coring sites are
selected on the basis of geomorphological descriptions.
Detailed three-dimensional pollution mapping is needed
along embanked rivers (Hobo et al. 2014), along rivers
with significant natural or artificial channel shifts
(Ciszewski and Malik 2004; Matys Grygar et al. 2013;
Ciszewski et al. 2014; Foulds et al. 2014), and
particularly on floodplains with a temporally variable
aggradation/erosion balance (Macklin et al. 1994;
Miller et al. 1998). In floodplains with a more complex
microtopography and pollution history, heavy metals
may be concentrated in hotspots, the position of which
is hardly predictable (Heaven et al. 2000; Miller and
Orbock- Miller 2007; Matys Grygar et al. 2014; Fig. 4).
Extensive sampling is also required in catchments with
variable geology and geomorphology (Amorosi and
Sammartino 2007; Peh et al. 2008; Amorosi et al.
2014; Matys Grygar et al. 2016B) and numerous
pollution sources along the river course (Matys Grygar et al.
A faithful description of historically polluted
floodplains may require sampling the entire thickness of
floodplain fines (Lecce and Pavlowsky 2014; Matys
Grygar et al. 2013, 2014, 2016a). The average vertical
deposition rate of overbank fines by a medium-sized
river is of the order of millimetre per year, and
inchannel sedimentation, including oxbow lakes, may be
up to a few centimetre per year (Sedláček et al. 2016).
Century-old polluted strata can hence be expected at
depths between a few centimetre to a few metre, the
latter being particularly probable in the channel belt
(Matys Grygar et al. 2013, 2016b). If erosion basis
and/or floodplain level changed in or since the pollution
period, each geoform in floodplain should be examined
(Macklin et al. 1994; Miller et al. 1998). To describe the
Fig. 4 Distribution of U in top 0–
10 cm in floodplain of the
Ploučnice River, typical for rivers
dominated by metal pollution of
fine grained sediments mixed
with unpolluted parent coarser
grained sediment. Data obtained
by in situ (portable, handheld)
XRF mapping. Lower
concentrations of U (<90 ppm)
are in coarser sediments in areas
covered by less polluted
postmining sediments in proximal
floodplain, higher concentrations
are in finer distal floodplain
deposits and in certain parts of
levee (J. Elznicová, unpublished
internal floodplain architecture, Houben (2007)
recommended a series of drill cores across the floodplain with
spacing less than the current channel width. In polluted
floodplains, that approach should be applied at least to
the range of channel shifts during the pollution period
and the nearby floodplain (Notebaert et al., 2011; Matys
Grygar et al. 2013). Because the thickness of polluted
strata can also vary for a single flood deposit from the
order of mm to the massive, decimetres thick laterally
deposited body, sampling in cores should be continuous
with steps in the order of cm.
Any analytical method can be used for pollution
mapping if it is sufficiently productive. AAS, ICP MS
or ICP OES and X-ray fluorescence spectroscopy
(XRF) are currently the most popular methods,
employed by many commercial laboratories. To
enhance the analytical productivity at low costs, there is
a tendency to replace methods that require time- and
cost-demanding dissolution steps (total decomposition
with mixed acid digestion or melting) by XRF with
pressing or only pouring samples to measuring cells
(Matys Grygar et al. 2014; Perroy et al. 2014). To get
around the total decomposition, pseudototal decay
(aqua regia or single acids) has been proposed and is
still used. However, it is not suitable to quantify most
lithoge nic elements s uitable f or geoche mical
Several types of portable instruments are suitable for
direct pollution mapping (Gałuszka et al. 2015; Horta et
al. 2015), but they are primarily tested for soils and
industrial, mining and waste disposal sites. Portable
XRF (PXRF) spectrometers are particularly promising
for pollution studies: In situ analyses require no
sampling/sample pre-treatment and can easily produce
pollution maps (Hürkamp et al. 2009), identify pollution
hotspots (Weindorf et al. 2012) and instantaneously
produce information essential for decision making in
technological operations (Lemiere et al. 2014). The
determination limits of PXRF are usually sufficient for
elements such as Pb and Zn, for elements such as As, Cu
and Ni in moderately polluted sediments, and for most
risk elements in severely polluted sites. Varying
humidity, organic matter content and texture effects may be a
hindrance, the handling of which requires empirical
corrections and further work (Lemiere et al. 2014;
Weindorf et al. 2014).
Other methods for direct in situ mapping are less
common. For radioactive pollution, gamma-activity
mapping should be the method of the first choice
(Martin et al. 2015). In situ gamma spectrometry can
also produce lithological mapping in floodplains
(Spadoni and Voltaggio 2013). Attempts have been
made to use Vis-NIR spectrometers (VNIR) with
empirical calibration for soil- or floodplain-pollution
mapping. The expectations have been based on the
successful correlation of heavy-metal concentrations with the
concentration of their main carriers, Fe-oxides, organic
matter and clay minerals obtained by laboratory VNIR
measurements (Kooistra et al. 2001, 2003). The
hindrance for in situ mapping is still the problem of varying
humidity and the variable state of the vegetation in the
Floods play a crucial role in the remobilization of heavy
metals from historically polluted deposits, whereas
present-day pollutants are primarily transported during
moderate and low flows. Moreover, floods, which
normally represent only a small percentage of annual
discharge, are a phenomenon that creates and reshapes the
floodplain and is responsible for the transfer of metal
pollutants from temporary sinks in the channel. In
perennial rivers, metals in dissolved forms or associated
with fine sediments can be transported over long
distances depending on the flow competency, whereas
sediments can be more persistently stored in a floodplain
at overbank flows. For this reason, the length of the river
and the degree of its modification, including structures
such as groynes, weirs, dams, oxbow lakes or side
channels, which are important traps for fine sediments,
control the pollutant conveyance through the river
system. The river damming and variety of engineering
measures in floodplains suppress the effect of floods
and limit overbank processes in most rivers.
Floods control the distribution of
sedimentassociated heavy metals in alluvial channels by the
creation of mezoforms, which are then modified during
moderate and low flows and filled or infiltrated with
polluted fines. Sediment sorting in a channel during
both floods and lower discharges and in a floodplain
during floods is a highly important mechanism
responsible for rapid changes in heavy-metal concentrations in
sedimentary sequences. The variable sorting overlaps
the record of the historical changes of river pollution,
and all reconstructions of heavy-metal pollution from
those sequences must correct concentrations for
Proportions of flood-related physical and chemical
metal mobilization from historically contaminated
floodplains to rivers are hardly quantifiable because they
are so site specific. However, both processes are most
intensive during certain stages of floods. Typically, the
physical remobilization of metals seems to dominate
over leaching under high discharges in reaches where
fine-grained alluvia are easily eroded. By contrast,
floodplains constitute diffusive metal sources even in
laterally stable reaches, particularly if polluted fines
occur in a framework of coarse-grained sediments
favouring the migration of dissolved metals. The
relative contribution of physical and chemical metal
mobilization depends not only on the rate of river erosion and
the average grain size of alluvia but also on the depth
and frequency of water-table fluctuations.
Flood-related metal storage and remobilization are
controlled by river channelization, but their influence
depends on the timing and extent of the engineering
works. Generally, the accretion of polluted sediments
in groyne basins and in cut-off channels, performed not
long before the pollution period, makes them hotspots of
pollutants transported in enhanced amounts during
high-water stages. Moreover, floodplains of lined river
channels that adjust to new, post-channelization
hydraulic conditions become the permanent sink for fine and
polluted sediments, which accumulate solely during
overbank flows. In such laterally stable river reaches,
metal mobilization occurs only by slow leaching, and
their sediments, which accrete at a moderate rate, are the
best archives of the catchment pollution with heavy
Acknowledgments The work by TMG was supported by the
Czech Science Foundation (project 15-00340S). Z. Hájková (IIC
Řež) is thanked for technical help in the preparation of the
manuscript. J. Elznicová (J.E. Purkyně University in Ústí nad Labem,
Czech Republic kindly provided a figure (pollution mapping).
Open Access This article is distributed under the terms of the
Creative Commons Attribution 4.0 International License (http://
creativecommons.org/licenses/by/4.0/), which permits
unrestricted use, distribution, and reproduction in any medium, provided
you give appropriate credit to the original author(s) and the source,
provide a link to the Creative Commons license, and indicate if
changes were made.
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